Optical properties of chromophoric dissolved organic matter (CDOM) and dissolved organic carbon (DOC) levels in constructed water treatment wetland systems in southern California, USA

Journal Pre-proof Optical properties of chromophoric dissolved organic matter (CDOM) and dissolved organic carbon (DOC) levels in constructed water treatment wetland systems in southern California, USA Catherine D. Clark, Warren J. De Bruyn, Benjamin Brahm, Paige Aiona PII:

S0045-6535(20)30098-9

DOI:

https://doi.org/10.1016/j.chemosphere.2020.125906

Reference:

CHEM 125906

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Received Date: 9 September 2019 Revised Date:

7 December 2019

Accepted Date: 12 January 2020

Please cite this article as: Clark, C.D., De Bruyn, W.J., Brahm, B., Aiona, P., Optical properties of chromophoric dissolved organic matter (CDOM) and dissolved organic carbon (DOC) levels in constructed water treatment wetland systems in southern California, USA, Chemosphere (2020), doi: https://doi.org/10.1016/j.chemosphere.2020.125906. This is a PDF file of an article that has undergone enhancements after acceptance, such as the addition of a cover page and metadata, and formatting for readability, but it is not yet the definitive version of record. This version will undergo additional copyediting, typesetting and review before it is published in its final form, but we are providing this version to give early visibility of the article. Please note that, during the production process, errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain. © 2020 Published by Elsevier Ltd.

Optical properties of chromophoric dissolved organic matter (CDOM) and dissolved organic carbon (DOC) levels in constructed water treatment wetland systems in Southern California, USA.

Catherine D. Clark*a, Warren J. De Bruynb, Benjamin Brahmb and Paige Aionab

a. Department of Chemistry, Western Washington University, 516 High Street, Bellingham, WA 98229, USA b. Schmid College of Science and Technology. Chapman University, One University Drive, Orange, CA 92780, USA

*Corresponding author: [email protected] 714-808-3227

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Optical properties of chromophoric dissolved organic matter (CDOM) and dissolved

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organic carbon (DOC) levels in constructed water treatment wetland systems in Southern

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California, USA.

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Catherine D. Clark*a, Warren J. De Bruynb, Benjamin Brahmb and Paige Aionab

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a. Department of Chemistry, Western Washington University, 516 High Street, Bellingham, WA 98229, USA b. Schmid College of Science and Technology. Chapman University, One University Drive, Orange, CA 92780, USA

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*Corresponding author:

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[email protected]

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714-808-3227

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Abstract

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Many removal mechanisms in treatment wetlands involve absorption to organic matter. Optical

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properties and DOC levels of waters entering and exiting 4 treatment wetland systems in Orange

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County, Southern California, were measured to characterize the dissolved organic matter pool.

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Average DOC levels decreased between the inlets and outlets, except for Forge Street (FS),

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which increased. For 3 wetlands, absorption coefficients decreased between inlet and outlet; the

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exception was FS, which increased. Average spectral slopes for the inlets and outlets were

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similar. Average intensities of terrestrial humic peaks A and C from 3D EEM fluorescence

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spectra decreased between the inlets and outlets for most wetlands. No EEM protein peaks were

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observed. Average flu/abs ratios for inlets and outlets were similar (high point for FS inlet

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excluded), suggesting chromophoric dissolved organic matter (CDOM) of a similar composition

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was present. The average FI value for the inlets and outlets was ~1.5, consistent with terrestrial

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sources of CDOM. Average BIX values for the inlets and outlets were ~0.8, suggesting limited

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contributions from autochthonous production of CDOM. Dominant plant species in the wetlands

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were cattail and bulrush. Humic peaks A and C, along with protein peaks, were observed in plant

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leachates. Protein peaks rapidly degraded with solar simulator irradiation. Results indicate that

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most of the wetlands are a net sink for CDOM, possibly due to absorption to sediments. The FS

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wetland appears to have a source of non-CDOM optically active organic carbon, possibly from a

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pollutant.

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Keywords: CDOM; fluorescence; EEMs; wetland; water treatment; DOC.

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1. Introduction Wetlands can be used to treat various types of wastewater, including municipal waste

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waters, acid mine drainage, agricultural runoff and stormwater runoff (Cole, 1998). The use of

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wetlands for wastewater treatment was first investigated in the 1950s in Germany (Benezowsky,

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1996). In the 1980s and 1990s, in Europe, the United Kingdom and the USA, various wastewater

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treatment entities began using constructed wetland systems for wastewater treatment

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(Benezowsky, 1996). Constructed wetlands are engineered systems designed and built to use the

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processes that occur in natural wetland systems to treat waters within a more controlled

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environment (Vymazal, 2011). Constructed wetlands may be categorized based on hydrology,

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type of plant growth and flow path. Hydrology may be surface water flow or subsurface flow,

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vegetation may be free floating, emergent or submerged and the flow path may be horizontal or

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vertical (Vymazal, 2008). Examples of constructed wetland systems across the world are given

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in Shutes (2001) and in various states in the USA in Carlisle and George (1991). Surface flow

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systems with emergent vegetation that resemble natural marshes are most common in North

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America (Cole, 1998); the flow path through these systems is horizontal. Cattail and bulrush are

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emergent plant species used in many treatment wetland systems in the USA (Vymazal, 2008).

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Some studies have explored the use of native plant species like prairie cordgrass in treatment

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wetlands (Bonilla-warford and Zedler, 2002).

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Surface flow treatment wetlands remove contaminants through a variety of mechanisms

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(Vyzamal, 2008). Suspended solids are removed through physical processes like sedimentation,

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aggregation and surface adhesion. Soluble organic compounds can be degraded aerobically and

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anaerobically by microbial decomposition. Wetlands with emergent vegeatation typically have

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aerated zones near the water surface because of the diffusion of oxygen into the water column, 3

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and anoxic zones in and near the sediments. Nitrogen is removed by nitrification/denitrification

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processes. Ammonia is oxidized by nitrifying bacteria in aerobic surface waters, and nitrates are

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removed in anoxic zones by denitrifying bacteria. Nitrogen can also be removed by vegetation

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growth, but this loss is usually negligible compared to inputs. Phosphorus is removed through

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adsorption to sediments or organic matter, complexation, precipitation, microbial assimilation or

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plant uptake. Metals are removed by precipitation, plant growth and adsorption to organic matter

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and sediments. Other species including chemicals and pathogenic organisms may be removed by

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a combination of physical, chemical and biological processes. These include oxidation, exposure

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to ultraviolet radiation and absorption by organic matter. Several studies on constructed wetlands

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have shown the removal of nutrients (Kao and Wu, 2001; Adhikari et al., 2011), metals (Nelson

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et al., 2006; Adhikari et al., 2011; Haarstad et al., 2012), various organic pollutants (Haarstad et

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al., 2012) and pharmaceuticals (Lee et al., 2014; Prasse et al., 2015).

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Many of the removal mechanisms described above involve absorption to organic matter.

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Another mechanism for removal by organic matter is through indirect photoxidation by reactive

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intermediates like singlet oxygen and hydroxyl radicals produced when optically active dissolved

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organic matter in the water column absorbs sunlight (Sardana et al., 2011). Understanding the

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nature and cycling of organic matter in treatment wetlands is thus important in the context of

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understanding the dynamics of pollutant processing in these systems. Dissolved organic matter

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(DOM) in the water column in natural waters is a heterogeneous mixture of large water-soluble

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organic molecules produced from the breakdown of terrestrial plant litter material or from

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microbial processes (McKnight and Aiken, 1998; Zsolnay et al., 1998). The quality of DOM

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from various sources affects its transformation and degradation during photochemical and

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microbial processes (Moran and Zepp, 1997; Moran et al., 2000; Obernosterer and Benner, 4

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2004). Chromophoric dissolved organic matter (CDOM) is the sunlight-absorbing fraction of

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DOM that mediates photochemical processes in aquatic systems (Green and Blough, 1994;

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Coble, 1996). Optical properties of CDOM measured by absorbance and fluorescence

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spectroscopy have been used to assess the sources and structure of DOM (see for example Coble,

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1996; McKnight and Aiken, 1998). For example, spectral slopes of the absorbance spectrum

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have been used to estimate molecular weight and aromatic content (Belize and Guo, 2006;

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Helms et al., 2008), with DOM from terrestrial plants typically having more aromatic and

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conjugated structures leading to absorbance at longer wavelengths than microbially-derived

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DOM, and hence lower spectral slopes (del Vecchio and Blough, 2002).

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We have previously used absorbance and fluorescence spectroscopy to characterize the

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optical properties of CDOM in the surface and pore waters of salt marshes in this region (Clark

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et al., 2008 and 2014; Bowen et al., 2017). Absorbance spectroscopy was used as a tool to

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characterize DOM in a restored urban marsh (Elbishilawi and Jaffe, 2015) and in treatment

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wetlands (Barber et al., 2001). Three-dimensional excitation-emission matrix (EEM)

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fluorescence spectroscopy has been previously used to examine the changes and characteristics

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of DOM in constructed wetland systems (Du et al., 2014; Yao et al., 2016; Sardana et al., 2019).

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There is some discrepancy in the literature as to whether constructed wetlands are a source or

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sink for dissolved organic carbon (DOC). Some studies have shown a decrease in DOC levels

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across constructed wetlands receiving wastewaters (Pinney et al., 2000) and runoff (Stern et al.,

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2007), whereas others have shown an increase in DOC levels across treatment wetlands

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receiving wastewaters (Barber et al., 2001) and in wetland mesocosms modelling wetland

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treatment systems for runoff (Villa et al., 2014).

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In this study, we measured DOC levels and optical properties of CDOM at the inflow and

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outflow of 4 constructed treatment wetland systems for dry weather runoff in Orange County,

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Southern California, to assess whether these systems were a source or sink for carbon and how

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the properties and characteristics of the DOM pool changed as it traversed the wetland.

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2. Methods

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2.1 Sampling. The Irvine Ranch water District (IRWD) was established 58 years ago in Orange

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County, Southern California, USA to manage water supply and water treatment in the region. In

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2005, the IRWD began construction on a selection of a planned total of 42 man-made wetland

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systems. These wetlands are designed for the treatment of dry and low flow urban runoff, and

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smaller storm water runoff, in the San Diego Creek/Newport Bay watershed. Water is diverted

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into the man-made wetlands in the absence of major storm events. The wetlands remove 67% of

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nitrogen, 74% of orthophosphate and 71% of bacteria, along with other contaminants like

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selenium (60% removal) and sediments (90% removal)(Irvine Ranch Water District NTS fact

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sheet: https://www.irwd.com/images/pdf/facilities/nts/nts_fact_sheet.pdf). Each site varies in

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total surface area from approximately 2,000 to 20,000 m2, in addition to an older well-

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established main treatment plant that contains over 180,000 m2 of natural treatment wetlands. In

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total, the natural treatment wetland sites (NTS) cover 180 acres and drain 15,000 acres, around

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20% of the watershed area. These NTS are freewater surface constructed systems with emergent

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macrophytes (Vymazal, 2008). The processes used to purify water in the artificial wetland

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system are dominated by 2 plant species, both native to natural southern Californian freshwater

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systems, cattail (Typha latifolia) and bulrush (Scirpus atrovirens). Four sites were chosen to

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sample in June 2009 based on proximity to the laboratory: Forge Street (FS; N 33°73.493′, W

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117°76.273′, ≈8,800 m2, 23 June), Quail Springs (QS; N 33°65.571′, W 117°78.079′, ≈28,000

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m2, 16 June), Red Hill (RH; N 33°72.206′, W 117°82.550′, ≈1950 m2, 22 June), and Turtle Ridge

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(TR; N 33°63.335′, W 117°82.384′, ≈18,000 m2, 15 June). The sampling locations are shown in

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Figure 1. The residence time for the largest system, Quail Springs, was given as being on the

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order of 10 days in one study (Wyss, 2017). Residence times for the other wetland sites have not

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been reported. One sample of 500 mL was taken from surface waters at both the inlet and outlet

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of each site between 10 am and 12 noon on the day sampled. At the time of sampling, these were

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relatively new wetland systems, less than 4 years old. These sites eventually drain into the San

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Joaquin channel and then into the Upper Newport Back Bay, an estuary and ecological reserve,

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which empties into the Pacific Ocean.

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Fig 1 Schematic of the 4 treatment wetland sampling locations in Orange County, California,

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USA.

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2.2 Measurements. After collection of the surface water samples, they were immediately tested

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for temperature and pH using a hand-held probe (Hanna Instruments; HI9828). Once back in the

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laboratory, they were vacuum filtered through 0.2 µm filters (Whatman, Durapore), then placed

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in covered containers and refrigerated at 4 C˚ until further analysis. The total organic carbon

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contained in each surface water sample was measured using a Shimadzu TOC analyzer, with a 3

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minute sparge time, a 2% HCl acidification solution, and a deionized water blank.

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Absorbance was measured using a diode-array UV-visible spectrometer (Agilent 8453).

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Spectra were obtained in a quartz cell with a path length of 1 cm using nanopure water as a

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blank. The wavelength range was from 200-700 nm. The spectral resolution of the instrument is

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<2 nm and stray light is <0.03%. The absorption coefficient (in units of m-1) was calculated from

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the measured absorbance using Eq. (1) (Hu et al. 2002):

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α = [(2.303Abs)/(l)]

(1)

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where Abs is the absorbance, α is the absorption coefficient, and l is the path length of the quartz

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cell in m (0.01 m in this study).

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Spectral slopes (S) were calculated from the absorbance spectra for a wavelength range of 300-

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400 nm from Eq. (2) (Helms et al. 2008):

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− =



αλ α

λ λ

(2)

where αλ is the absorption coefficient at wavelength λ and αo is the absorbance at a reference wavelength λ0 (Green and Blough 1994; Moran et al. 2000).

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A scanning fluorometer (Quantamaster, PTI) was used to measure fluorescence spectra.

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Three-dimensional excitation-emission matrix fluorescence spectra (EEMs) were obtained by

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ranging excitation wavelengths from 260-430 nm in 5 nm increments and emission wavelengths

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from 260-650 nm, also in 5 nm increments. A nanopure water EEM was obtained daily in order

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to subtract out the Raman peak of water and the Rayleigh scattering. Spectra were corrected for

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instrumental response using the correction file from the manufacturer (based on factory

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calibration). The % error on 3 duplicate absorbance and fluorescence scans was <0.5%.

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Fluorescence intensities were generated by the instrument in units of photons s-1. These units

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were converted to QSU (quinine sulfate units, where 1 QSU = 1 ppb quinine sulfate in 0.05 M

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H2SO4) using a calibration curve (Mopper and Scultz, 1993). A mass of 100 mg of quinine

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sulfate was dissolved in 1 L of H2SO4 with a concentration of 0.1 N to create the standard QSU

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solution.

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The fluorescence index (FI; McKnight et al., 2001) and the index of recent autochthonous

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contribution (BIX; Huguet et al., 2009) were calculated from the EEM spectra. FI, the ratio of

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the emission intensities at 450 to 500 nm for an excitation wavelength of 370 nm, is used to

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identify CDOM arising from microbial and terrestrial sources. BIX is used to identify CDOM

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produced from recent biological activity; this index is calculated at an excitation wavelength of

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310 nm from the ratio of the emission intensities at wavelengths of 380 to 430 nm.

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2.3 Photodegradation experiments. Senescent plant material samples of cattail (Typha latifolia)

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and bulrush (Scirpus atrovirens) were retrieved from the Turtle Ridge site. They were soaked in

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DI water for 24 hours under laboratory conditions in the dark at room temperature, then filtered

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through 0.2 µm filters. Portions of the sample were then irradiated in a solar simulator (Luzchem

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Solsim) for up to 3 hours. The solar simulator uses a 300 W xenon lamp to reproduce an AM 1.5

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solar spectrum. The total photon flux measured at the time of these irradiations by nitrite/benzoic

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acid/hydroxybenzoic acid chemical actinometery (Jankowski et al., 1999) was 3.7 x 1019 photons

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m-2 s-1. The lamp is positioned approximately 8 inches away from the sample and the simulator

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chamber is flushed continually with air during the irradiation to control temperature. Optical

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properties of the sample were measured both before the irradiation as a control and then again

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immediately after irradiation.

Site

Abs coeff (300 nm)

Temp o C

pH

DOC (ppm)

spectral slope

peak A

peak C

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3. Results

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3.1 Water quality parameters Temperature for the 4 treatment wetland systems ranged from a

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low of 20.8 oC at the inlet of QS to a high of 26.9 oC at the FS inlet. Temperature increased by 4

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o

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of TR and FS by about 1.5 oC. The inlet and the outlet temperatures for the RH system were very

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similar (26.3 vs 26.5 oC). Overall the average inlet temperature (24.1±3.0 oC; 1σ; n=4) across all

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four sites was statistically the same (student’s t;-test 95%) as the average outlet temperature

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(24.3±2.4 oC 1σ; n=4) (Table 1). The pH ranged from a low of 7.22 for the TR inlet to a high of

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8.51 for the RH inlet. The pH increased by 0.14 between the TR inlet and outlet but decreased

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between the inlet and outlet for the other 3 wetland systems, ranging from a decrease of 0.09 for

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QS to a decrease of 0.83 for FS. The average inlet and outlet pH was 7.98±0.58 (1σ; n=4) and

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7.68±0.32 (1σ; n=4) respectively (Table 1). These are also statistically the same at the 95%

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confidence level (student’s t-test). The values of pH and temperature would be expected to

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change depending on the time of day sampled (due to changing solar radiation or algal activity

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for example), but the range of values reported here helps characterize these types of systems in

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this region in the summer.

C between the inlet and outlet of QS, but the temperature decreased between the inlet and outlet

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10

QS inlet

68.6

20.8

7.85

78.4

0.0188

409

330

QS outlet

41.5

24.7

7.76

44.2

0.0141

370

279

TR inlet

32.7

22.5

7.22

49.3

0.0111

266

188

TR outlet

29.2

20.9

7.36

32.7

0.0117

274

201

FS inlet

55.7

26.9

8.33

40.1

0.0140

546

494

FS outlet

76.5

25.2

7.5

56.0

0.0149

440

351

RH inlet

48.1

26.3

8.51

66.2

0.0130

437

298

RH outlet

42.8

26.5

8.09

43.9

0.0132

396

270

Inlet average ± 1σ σ

51.3±15.0

24.1±3.0

7.98±0.58

58.5±17.1

0.014±0.003

47.5±20.3

24.3±2.4

7.68±0.32

44.2±9.5

0.013±0.001

Outlet average ± 1σ σ

414±115

327±127

370±70

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Table 1. Water quality and optical parameters for the natural treatment wetland systems. The

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fluorescence intensities for peak A and C are given in QSU units.

275±61

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3.2 DOC measurements DOC values ranged from a low of 32.7 ppm for the TR outlet to a high

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of 78.4 for the QS inlet. In general, the DOC levels decreased between the inlets and outlets by

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20-30 ppm; the exception was the FS system, where DOC increased by 16 ppm between the inlet

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and outlet. The average DOC levels for the inlets and outlets were 58.5±17.1 (1σ; n=4) and

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44.2±9.5 (1σ; n=4) ppm respectively; also statistically identical (Table 1). These values are high,

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but this is not unusual for constructed wetlands. Measurements of DOC levels at the output of a

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treatment wetland in the Florida Everglades region ranged from 14 to 64 mg C/L (Stern et al.,

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2007). DOC values of 30-47 mg/L were measured at the outflow of mesocosms modelling

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treatment wetlands in the same region (Villa et al., 2014). In a study of treatment wetlands in

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Oregon and California, the DOC levels at the outflows reached a maximum of 29 mg C/L

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(Barber et al., 2001).

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3.3 Absorption measurements The absorption coefficients ranged from a low of 29.2 m-1 for the

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TR outlet to a high of 76.5 m-1 at 300 nm for the FS outlet. For 3 of the wetlands, the absorption

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coefficient decreased between the inlet and outlet, ranging from a decrease of 3.5 to 27.1 m-1.

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The exception was the FS system, which increased by 21 between the inlet and outlet. Average

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absorbance coefficients for the inlets (51.3±15.0 ; 1σ; n=4) and outlets (47.5±20.3 m-1 ;1σ; n=4)

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were statistically the same at the 95% confidence level (Table 1). Figure 2 is a plot of the

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absorption coefficients at 300 nm vs. 254 nm for all 4 of the treatment wetlands. The line shown

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is from a linear regression with R2=0.98.

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The spectral slope ranged from a high of 0.0188 for the QS inlet to a low of 0.0111 for

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the TR inlet. For 3 of the wetland systems, the spectral slope increased slightly from inlet to

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outlet; for example, for TR, the value increased from 0.0111 to 0.0117. The exception was the

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QS system, which decreased from 0.0188 at the inlet to 0.0141 at the outlet. The average spectral

235

slopes for the inlets and outlets was 0.014±0.003 (1σ; n=4) and 0.013±0.001 (1σ; n=4)

236

respectively (Table 1); once again statistically identical. Typically, higher spectral slope values

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indicate low molecular weight material and/or decreasing aromaticity (Blough and Del Vecchio,

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2002; Helms et al., 2008).

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Fig 2 Absorbance coefficient at 300 nm vs. absorbance coefficient at 254 nm for all natural

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treatment wetland samples (□ inlet; ○ outlet). Line shown is from a linear fit. Slope = 0.59;

242

intercept = -2.84; R2 = 0.98

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The SUVA ratio was calculated from the absorption coefficient at 254 nm divided by the

245

concentration of DOC in ppm (Hansen et al., 2016). Results are given in Table 2. SUVA values

246

represent absorbance per unit carbon. Typically, lower values are associated with lower aromatic

247

content (Weishaar et al., 2003). Values ranged from a low of 1.06 for the TR inlet to a high of

248

3.37 for the FS inlet. In general, values increased from inlet to outlet, except for the FS system

249

where values decreased. The value for the FS inlet at 3.37 is high, with the other inlets ranging

250

from 1.06 to 1.45. If we treat the FS point as an outlier, the average for the inlets without the FS

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point is 1.30±0.21 (1σ; n=3). The average for the outlets is 1.76±0.08 (1σ; n=4). Without the FS

252

point, these averages are statistically different at the 95% confidence level (student’s t-test).

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Site QS inlet

SUVA 1.45

Flu/abs 4.7

FI 1.47

BIX 0.70

QS outlet

1.69

5.8

1.56

0.80

TR inlet

1.06

4.9

1.53

0.80

TR outlet

1.77

6.2

1.56

0.79

FS inlet

3.37

7.9

1.51

0.74

FS outlet

1.77

4.0

1.52

0.77

RH inlet

1.38

4.8

1.49

0.78

RH outlet

1.87

5.0

1.50

0.83

Inlet average ± 1σ σ

1.30±0.21a

4.8±0.08b

1.50±0.02

0.76±0.04

Outlet average ± 1σ σ

1.76±0.08

5.2±1.0

1.53±0.03

0.80±0.02

254

a: excluding the FS inlet point; with this point included, 1.81±1.05

255

b: excluding the FS inlet point; with this point included, 5.6±1.3

256

Table 2. SUVA values and fluorescence indices for the natural treatment wetland systems.

257

258

3.4 Fluorescence measurements The 3D excitation-emission matrices (EEMs) were used to

259

identify specific CDOM sources (Coble, 1996; de Souza Sierra et al. 1997; McKnight et al.

260

2001). The major peaks identified by EEMS are shown in Table 3. These are: 1) peak A,

261

terrestrial humic-like; 2) peak T, tryptophan protein-like; 3) peak B, tyrosine protein-like; 4)

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262

peak M, marine humic-like (associated with microbially produced material); and 5) peak C,

263

terrestrial humic-like.

264

Peak

Exmax (nm)

Emmax (nm)

Specified Peak (Ex/Em)

Component type

C A M B T

330 – 350 240 – 260 310 – 320 270 – 280 270 – 280

420 – 480 380 – 460 380 – 420 300 – 320 320 – 350

340/435 260/450 -275/300 275/350

Terrestrial, humic-like Terrestrial, humic-like Marine, humic-like Tyrosine, protein-like Tryptophan, protein-like

Table 3. Major fluorescence bands for natural water EEMs (Coble, 1996).

265

266

The fluorescence intensity for peak A ranged from a high of 546 QSU for the FS inlet to

267

a low of 266 QSU for the TR inlet. The same trend was observed for the fluorescence intensity

268

for peak C, which ranged from a high of 494 QSU for the FS inlet to a low of 188 QSU for the

269

TR inlet. The fluorescence intensity of peak A and C decreased for 3 of the wetland systems

270

between the inlet and outlet; the exception was the TR system where the intensities increased

271

from inlet to outlet. DOC levels decreased between inlet and outlet for the same wetland. DOC is

272

a measure of the dissolved organic carbon levels, whereas the fluorescence intensities are a

273

measure of the fluorescent fraction of the DOM. These are often correlated with each other, but

274

they are not necessarily a measure of the same substances. It is conceivable that the water

275

entering the TR wetland may have had a source of dissolved carbon that was not fluorescently

276

active in the regions where peak A and C fluoresce.

277 278

The average intensities of peak A for the inlets and outlets was 414±115 (1σ; n=4) and 370±370 (1σ; n=4) respectively; for peak C, the average intensity for the inlets was 327±127

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279

(1σ; n=4) and 275±61 (1σ; n=4) for the outlets (Table 1). Figure 3 shows EEMs for the inlet and

280

outlet of the Quail Springs natural treatment wetland system. These are typical of the EEMs

281

measured for the study sites. The terrestrial humic-type peaks A and C are observable in both

282

EEMs, with reduced intensities in the outlet EEM. There are no protein peaks observed; these are

283

typically attributed to fresh production occurring.

284

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285 286

Fig 3 EEMs for the inlet and outlet of the Quail Springs treatment wetland system. Intensity

287

scale is in QSU units

288

289

The fluorescence indices are given in Table 3. The ratio of the fluorescence at an

290

excitation of 350 nm and an emission of 450 nm to the absorption coefficient at 300 nm is called

291

the flu/abs value. The flu:abs ratio is a measure of CDOM composition changes, with decreasing

292

values associated with increasing molecular weight, humification and aromaticity (Belzile and

293

Guo, 2006). This ratio ranged from a high of 7.9 for the FS inlet to a low of 4.0 for the FS outlet.

294

The value increased from the inlet to outlet for 3 of the wetland systems, but decreased for the

295

FS. The average flu/abs value for the inlets was 4.8±0.08 (1σ; n=3) with the high point for the FS

296

inlet excluded, while the average for the outlet was statistically similar (student’s t-test) at

297

5.2±1.0 (1σ; n=4). Figure 4 shows a plot of the fluorescence in QSU at an excitation of 350 nm

17

298

and emission of 450 nm vs. absorbance coefficient at 300 nm for all natural treatment wetland

299

samples.

300

SUVA values indicated that the aromatic content generally increased through the wetlands

301

as indicated by the increase between inlets and outlets in Table 2, but an opposite result was

302

obtained according to the flu:abs ratio. These two parameters measure two different things: the

303

amount of CDOM vs. DOC for the SUVA and the amount of FDOM vs. CDOM for the flu/abs

304

ratio. The SUVA ratio is likely more sensitive to changes in aromaticity than the flu:abs ratio.

305

It’s also worth noting that while the flu:abs ratio generally did increase from inlet to outlet for

306

most of the wetlands, which would indicate a decrease in aromaticity, the values are very similar

307

within the range measured. This is seen in Figure 4 plotting fluorescence vs. absorbance, where

308

all points except for 1 fall on a straight line.

309 310

Fig 4 Fluorescence in QSU at an excitation of 350 nm and emission of 450 nm vs. absorbance

311

coefficient at 300 nm for all natural treatment wetland samples (□ inlet; ○ outlet). Line shown is

18

312

from a linear fit. Slope = 3.19; intercept = 82.8; R2 = 0.87; linear fit excludes the outlying high

313

point for the FS inlet. The FI value ranged from a low of 1.47 for the QS inlet to a high of 1.56 for the QS and

314 315

TR outlets. The value for all 4 wetlands increased slightly from the inlet to the outlets. The

316

average value for the inlets was 1.50±0.02 (1σ; n=4) while the value for the outlet was

317

statistically identical (student’s t-test) at 1.53±0.03 (1σ; n=4). The FI value is an indicator of the

318

relative contribution of terrestrial and microbial sources to the CDOM pool (Hansen et al., 2016).

319

The f450/f500 value is about 1.9 for aquatic and microbial sources and about 1.4 for terrestrial and

320

soil sources (McKnight et al., 2001; Huguet et al., 2009). The BIX value ranged from a low of

321

0.70 for the QS inlet to a high of 0.83 for the RH outlet. Values increased slightly from inlet to

322

outlet for 3 of the wetlands, with the values decreasing slightly for the TR wetland. The average

323

value for the inlets was 0.76±0.04 (1σ; n=4), whereas the average value for the outlets was

324

0.80±0.02 (1σ; n=4). Once again, these averages are statistically identical. BIX values increase

325

as contributions from recently produced CDOM of autochthonous origin increase (Huguet et al.,

326

2009).

327

Figure 5 shows the EEMs obtained for the cattail plant leachate before and after 2 hours

328

of irradiation with a solar simulator. Both EEMs show the presence of humic-type peaks A and

329

C, with peak A at a higher intensity (138 QSU) than peak C (120 QSU). After irradiation, both

330

peaks A and C are reduced in intensity by approximately 20 QSU to 80% of their original values.

331

Both EEMs also show the presence of a protein peak, the tryptophan-like peak T, at a higher

332

intensity than peaks A and C (250 QSU pre-irradiation). This is also reduced in intensity after the

333

irradiation, to 152 QSU, or 60% of its initial value.

19

334

335 336

Fig 5 EEMs for the cattail plant leachate before (0 hours) and after irradiation with a solar

337

simulator (2 hours). Intensity scale is in QSU units

338

20

339

340

341 342

Fig 6 EEMs for the bulrush plant leachate before (0 hours) and after irradiation with a solar

343

simulator (2 hours). Intensity scale is in QSU units 21

344

Figure 6 for the bulrush plant leachate shows a similar trend, with the presence of both

345

peaks A and C at similar intensities (130 and 125 QSU respectively pre-irradiation), and reduced

346

intensities after irradiation to 78 and 88% of their original values respectively. Peak T is also

347

present in these EEMs, at a higher intensity than for the cattail prior to irradiation (427 QSU).

348

This peak is also reduced in intensity after irradiation to 200 QSU or 47% of its original value, as

349

for the cattail leachate. Irradiation for an additional hour did not change the fluorescence

350

intensities of the peaks that were observed after 2 hours of irradiation.

351

352

4. Discussion

353

Studying the optical properties of CDOM in natural water systems allows us to evaluate

354

the relative amounts and sources of this material. Changes in the fluorescence intensity and

355

absorption coefficient are related to changes in both the structure and levels of CDOM, whereas

356

changes in calculated parameters like the spectral slope and fluorescence indices are related to

357

changes in CDOM functionality only, for example aromaticity and relative molecular weight

358

(McKnight et al., 2001; Tzortziou et al., 2011). The composition and levels of DOM may change

359

in natural water systems due to differences in sources (terrestrial vs. microbial), ageing

360

processes, physical processes (eg. dilution, absorption to sediments), and photochemical and

361

biological processes.

362

4.1 Wetlands as a carbon sink The trends in absorbance coefficient and DOC concentrations are

363

consistent with each other, with 3 of the 4 wetlands showing a decrease in both parameters between the

364

inlet and the outlet (and 1 showing an increase). We had expected these parameters to increase through

365

the wetland due to the input of material produced by the wetland plants. This decrease suggests that the

22

366

organic carbon pool is being depleted as water flows through the wetland, possibly due to physical

367

absorption processes to sediments or biological and photochemical processing. This indicates material

368

being produced in situ from plants in the wetland is not enough to offset this sink in the organic carbon

369

pool. These wetlands were newly established within the last 4 years at the time of the study. This may

370

explain why carbon could be efficiently sequestered in the sediments as a possible sink, as there had not

371

been sufficient time to build up high sediment DOM levels from absorption processes.

372

The QS wetland had a decrease in DOC levels between inlet and outlet to 56% of the inlet value

373

The outlet values for the TR and RH wetlands were 66% of the inlet values. The FS system by contrast

374

had an increase to 130% of the inlet value at the outlet. A study in the Florida Everglades in fresh water

375

natural and constructed wetlands found that these systems contained a labile DOC pool with rapid

376

turnover times, and that they were a net sink for DOC (Stern et al., 2007). Pinney et al. (2000) also noted

377

DOC removal through constructed wetlands; this varied seasonally, from a maximum of 49% to a

378

minimum of 9%. This is consistent with the observations here that DOC decreased through the wetlands.

379

In contrast, Barber et al. (2001) found that wetlands receiving wastewater for treatment produced a net

380

increase in DOC; one of their study sites was in Arcata, California. Villa et al. (2014) found that

381

mesocosms replicating treatment wetland systems in the Florida Everglades drainage system were in most

382

cases a source for DOC, with a few systems showing a decrease in DOC levels between the inlet and

383

outlet. These increases in DOC through the wetlands were attributed to biomass, with autochthonous

384

production occurring from wetland plants.

385

S values increase as photochemical and oxidative processes take place, producing less aromatic,

386

lower molecular weight material, and decrease for CDOM with higher aromaticity and humification

387

(Green and Blough 1994; Vodacek et al. 1997; Boyd and Osburn 2004, Stabenau et al. 2004; Twardowski

388

et al., 2004; Tzortziou et al. 2007). We had expected S values to increase through the wetland due to

389

photochemical processing of the CDOM in the water. The lack of a difference between the inlet and outlet

390

sites for 3 of the treatment wetlands suggests that whatever photochemical processing is taking place as 23

391

water flows through the wetlands is not sufficient to change the spectral slopes of the CDOM. This

392

suggests that photochemical processing may not be a significant sink for the DOM entering the wetland.

393

The primary sink may instead be absorption to the sediments. The values of 0.013-0.014 observed here

394

fall within the range of values previously observed for the surface waters of a coastal salt marsh in the

395

same county in Southern California (0.0051-0.016; Bowen et al., 2017). For the QS wetland, the decrease

396

in spectral slope between inlet and outlet might suggest an input of more aromatic and humified CDOM is

397

occurring. However, the decrease in the fluorescence intensities of peaks A and C between the inlet and

398

outlet for this wetland is not consistent with an input of CDOM as the water flows through the system. In

399

general, 3 of the 4 wetlands show a decrease in the fluorescence intensities of peaks A and C through the

400

wetland (along with decreases in absorption coefficient and, DOC concentrations), suggesting that the

401

wetlands are a net sink for CDOM rather than a source. The decrease in fluorescence intensities between

402

the inlets and outlets observed in this study is consistent with work by Yao et al. (2016). They showed a

403

decrease in fluorescent DOM occurred through a constructed wetland.

404

405

4.2 DOM sources An absorption ratio, the ratio of two CDOM absorption coefficients at two

406

different wavelengths, has been used in prior studies to indicate differences in molecular weight,

407

humification degree and the sources of CDOM in natural waters (Miller 1998, Clark et al., 2008,

408

Li and Hur, 2017). For example, in one study, the authors used an absorption ratio to help

409

differentiate between CDOM of terrigenous origin and higher molecular weight in an upper

410

estuary vs. CDOM of a more marine origin and lower molecular weight in the lower estuary and

411

coastal waters (Lei et al., 2019). Higher molecular weight CDOM has higher absorption at longer

412

wavelengths, while lower molecular weight CDOM absorbs more strongly at lower wavelengths.

413

A decrease of the ratio of an absorption coefficient at 254 nm vs. one at 300 nm could then

414

indicate an increase in the molecular size of CDOM. For this study, the absorption coefficients at 300 24

415

vs. 254 nm for the 4 treatment wetlands showed a strong linear relationship (Fig 1), indicating the same

416

absorption ratio. This suggests that the absorbing material entering and leaving the wetland waters was

417

similar and from the same source, likely terrestrial material from plants.

418

The FI value of 1.5 consistently observed in this study for the inlet and outlet of the wetlands is

419

consistent with a terrestrial source of CDOM (McKnight et al., 2001; Huguet et al., 2009). The BIX

420

values of 0.7 to 0.8 observed in this study for the wetland inlets and outlets is consistent with

421

values of 0.7-0.8 measured in two prior studies in a salt marsh in this region (Clark et al., 2014;

422

Bowen et al., 2017). These values are consistent with an allocthonous source, i.e. a terrestrial

423

plant source, rather than an autochthonous freshly produced microbial source in the water.

424

The plant leachate studies showed that two common natural treatment wetland plant

425

species do produce EEMs characteristic of CDOM with humic-type peaks from senescent

426

material, suggesting that biological production is occurring in the wetlands and that common

427

plants found in the wetlands are a source of CDOM as they decompose. In a prior study on a salt

428

marsh in this region, we found that leachates of common salt marsh plant species produced

429

CDOM (Clark et al., 2008). Rossel et al. (2013) found that the leachate of the common rush, a

430

wetland plant species, produced DOM, which was photochemically and microbially decomposed

431

over long time scales to produce material characteristic of deep-sea DOM. DOM was also

432

produced from leachates of the common rush in an earlier study (Vahatlo and Wetzel, 2008).

433

Another study on plant species in the Florida Everglades found that they were a source of DOC

434

to exported waters (Villa et al., 2014). Maie et al. (2006) found that leachates of cattail (one of

435

the species in this study) produced DOC, and that polyphenols in the leachate could be an

436

important source of CDOM.

25

437

This wetland plant source may not be enough to offset the sinks in the wetlands that reduce

438

CDOM levels, for example absorption to the sediments. The fluorescence intensities of peaks A

439

and C decreased with irradiation by a solar simulator, by ~20% over 2 hours of exposure,

440

suggesting that photodegradation may be an efficient removal mechanism for CDOM produced

441

from plants in the wetland. Clark et al. (2008) showed that salt marsh plant leachate humic peaks

442

degraded on the timescale of hours with irradiation from a solar simulator. Vahatlo and Wetzel

443

(2008) showed that solar radiation decomposes up to 99% of optically active wetland-derived

444

DOM from the common rush over long enough time scales. Bano et al. (1998) showed that

445

irradiation increased the extent of bacterial decomposition of DOC in a natural fresh water

446

wetland in Georgia. The plant leachates in this study also produced protein type peaks

447

characteristic of fresh production but these are rapidly photodegraded by sunlight over a

448

timescale of hours, which may be why they are not seen in the wetland water outlet EEMs. 4.3

449

DOM composition changes The SUVA values in general increased between the inlets and

450

outlets,indicating that the aromatic content of the dissolved organic material increased as water moved

451

through the wetland systems. This may be due to the input of DOM with a higher aromatic content or to

452

the photochemical and biological processing of the more aliphatic components of the input material as it

453

traverses the wetland. In a laboratory study of standard aquatic humic substances, irradiation with a solar

454

simulator decreased SUVA values due to photochemical processing (Sharpless et al., 2014). The fact that

455

our values generally increased between inlets and outlets is an indication that photochemical processing

456

may not be a significant driver of CDOM compositional changes in these wetlands. Barber et al. (2001)

457

observed a transformation in DOM to become more aromatic as it traversed through treatment wetlands,

458

which they attributed to the input of terrestrially derived material from the wetland plants. Pinney et al.

459

(2000) also saw an increase in SUVA values across a constructed wetland receiving lagoon-treated

460

wastewater. The unusually high SUVA value for the FS inlet (2-3 times higher than the other systems)

461

indicates that this wetland might have a different source of absorbing organic carbon, other than CDOM, 26

462

possibly a pollutant. The SUVA value of the FS outlet is similar to that for the other systems, suggesting

463

that this species may be removed on transport through the wetland.

464

The ratio of fluorescence intensity to the absorption coefficient can also reflect changes in CDOM

465

composition (Rochelle-Newall and Fisher, 2002). Lower values are associated with higher molecular

466

weight material and increasing humification and aromaticity (Belzile and Guo, 2006). The flu/abs value

467

reflects the relative amount of fluorescent DOM (FDOM) in the CDOM reservoir in the water system (as

468

represented by the absorption coefficient). Lower values indicate less fluorescent material in the CDOM

469

pool, suggesting changes in the structure of the DOM. The flu/abs ratio in this study was the same across

470

the wetlands, as indicated by the straight line fit to the data in Figure 3. The exception to this trend is the

471

FS wetland, which has an unusually high flu/abs value at the inlet (the outlier point in Figure 3). This

472

wetland also has a high SUVA value at the inlet and shows an increase in both absorbance coefficient and

473

DOC between the inlet and outlet, unlike the other wetlands. Other parameters for this wetland (pH,

474

temperature, spectral slope) follow the general trends for the other wetlands. The EEM for the FS inlet

475

shows no unusual features, containing only peaks A and C, as for the other wetlands. The fluorescence

476

intensities of the humic peaks A and C decrease between the inlet and outlet for this wetland, as they do

477

for most of the other wetlands. This suggests that there is not a significant net input of CDOM into the

478

wetland as water flows through it; if this input occurred, this would have explained the increase in

479

absorbance and DOC. These results may instead indicate that there is another source of optically active

480

organic carbon entering this wetland, possibly a contaminant. A compound that contains aromatic

481

rings in its structure would absorb light and fluoresce. Substances that may be found in urban

482

runoff that have aromatic rings and are optically active include PAHs, PCBs, DDT and

483

polychlorinated dioxins.

484

485

5. Conclusions 27

486

Two common natural treatment wetland plant species (cattail and bulrush) produced

487

CDOM as they decomposed, acting as a CDOM source in the wetlands. However, data from

488

absorption coefficients, DOC concentrations and humic peak fluorescence intensities indicate

489

that most of the wetlands are a net sink for CDOM. This may be due to a combination of

490

photochemical and biological processing, in conjunction with physical processes like absorption

491

to the sediments. The consistency in spectral slopes between the inlet and outlet of the wetlands

492

is not consistent with significant photochemical processing occurring. The optical characteristics

493

and fluorescence indices indicate that CDOM entering and leaving the wetlands studied is of a

494

similar terrestrial source that does not have a significant contribution from freshly produced

495

microbial material. The lack of evidence for photochemical or biological processing suggests

496

that adsorption to the sediments is the primary removal mechanism for DOM. One wetland

497

appeared to have a different source of non-CDOM optically active organic carbon, possibly a

498

pollutant.

499

500

Acknowledgements

501

We thank the Irvine Ranch Water District for access to the sampling sites.

502

503

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Highlights

•

Cat tail and bulrush are CDOM sources in the wetlands

•

The wetlands are a net sink for CDOM based on DOC and optical properties

•

Fluorescence indices suggest CDOM entering and leaving the wetlands is terrestrial

•

There is no significant contribution from freshly produced microbial material

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Adsorption to the sediments may be the primary removal mechanism for DOM

Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests: